Wastewater treatment plant (WWTP) effluents represent the main source of pharmaceuticals and illicit drugs (PIDs) in river systems in the U.S. Moreover, pharmaceuticals and illicit drugs (PIDs) contamination due to unsheltered homelessness, as is characterized by tents or sleeping bags along rivers and other waterways, could be another pressuring factor, but it has been studied scarcely. This study investigated the occurrence of PIDs in effluent and unsheltered homelessness-impacted sites in river systems in the U.S. In addition, potential ecological risks posed by selected pharmaceuticals to aquatic ecosystems were assessed based on risk quotients (RQs). Analyses were performed using Liquid Chromatography-Tandem Mass Spectrometry (LC-MS/MS). Several PIDs were found to be ubiquitous in the effluent-impacted sites (e.g. lidocaine, metoprolol tartrate, diclofenac acid and methamphetamine). Results also showed that PIDs contamination is ubiquitous in unsheltered homelessness-impacted sites. For example, the methamphetamine concentrations in Santa Cruz River ranged from 83.5 ng L− 1 to 450 ng L− 1. Antibiotics (6.94 ng L− 1 to 626 ng L− 1) were also detected in higher concentrations in Santa Cruz River. Pharmaceuticals except sulfamethoxazole were detected at concentrations far below Predicted-no effect concentration (PNEC) values across all effluent and unsheltered homelessness-impacted river systems, indicating negligible (low) or moderate ecological risks to algae and crustaceans. However, relying solely on single-compound risk assessment might underestimate cumulative ecological risk.
Research Article
Occurrence and ecotoxicological risks of pharmaceuticals and illicit drugs in effluent and unsheltered homelessness-impacted river systems
https://doi.org/10.21203/rs.3.rs-7273934/v1
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River systems
WWTP
unsheltered homelessness
pharmaceuticals
illicit drugs
LC-MS/MS
ecological risks
In recent years, the presence of contaminants of emerging concern (CECs) in surface waters has become one of the most pressing global environmental challenges (Drdanová et al. 2025). A wide range of land use types, including agricultural, residential, industrial, and commercial, contribute to the release of CECs into aquatic ecosystems (Kolpin et al. 2004; Fairbairn et al. 2016; Monk et al. 2025). Numerous studies have linked land use patterns and wastewater treatment plant (WWTP) effluents to the spatial distribution of pharmaceuticals and illicit drugs (PIDs) in the environment (Shala and Foster 2010; Fairbairn et al. 2016; Hain et al. 2023). Incomplete removal of many PIDs during conventional wastewater treatment continues to result in their discharge into receiving surface waters (Gautam et al. 2014; Wilkinson et al. 2022; Monk et al. 2025).
Urban rivers and streams are also impacted by diffuse and poorly characterized sources of pollution, further degrading water quality (Du et al. 2020). For example, fexofenadine and metformin have been detected at concentrations of up to 2.61 µg L− 1 and 1.41 µg L− 1, respectively, in U.S. rivers (Wilkinson et al. 2022). Long-term exposure to pharmaceuticals in aquatic environments has been shown to adversely affect fish and other aquatic organisms (Yang et al. 2015; Manjarrés-López et al. 2023). In one study, fish species including Leptocottus armatus and Oncorhynchus tshawytscha, collected from two WWTP-impacted estuarine sites in Puget Sound, exhibited elevated tissue concentrations (cumulative) of PIDs and personal care products, levels high enough to alter behavior, endocrine function, metabolism, and other physiological processes (Meador et al. 2017).
The entry of selected PIDs into surface waters has been associated with ecological risks such as the reduction of biological community structure, declining fish populations, and the promotion of antibiotic resistance (Batt et al. 2016; Fairbairn et al. 2016; Sims et al. 2024). Understanding the sources and distribution of PIDs within a watershed requires comprehensive, regionally specific evaluations to interpret source-sink relationships (Zhou et al. 2022). Prior research has found higher PID concentrations in surface waters more heavily influenced by anthropogenic activity (Omar et al. 2018; Biswas et al. 2022; Santos et al. 2022).
A growing concern in urban watersheds is the contribution of PIDs and microbial contamination from tent encampments located along riparian corridors. These areas, often inhabited by individuals experiencing unsheltered homelessness, typically lack adequate sanitation infrastructure, leading to open defecation and improper disposal of personal care items directly into waterways (Doerschlag 2021; Pinongcos et al. 2022). For instance, Gerrity et al. (2022) reported elevated concentrations of heroin, acetylmorphine, amphetamine, cocaine, acetaminophen, ibuprofen, and naproxen in Las Vegas washes impacted by unsheltered encampments. Historically, PID monitoring has focused largely on WWTP effluent (Gerrity et al. 2011; Skees et al. 2018; Bishop et al. 2020), while the influence of encampments studied scarcely.
To better understand the relative influence of treated wastewater and encampment-based contamination on surface water quality, we conducted a regionally targeted investigation across multiple river systems in the Western United States. We hypothesized that PID concentrations would be elevated in urban rivers impacted by WWTP effluent and that sites influenced by unsheltered homelessness would show disproportionately higher levels of pharmaceuticals, drugs of abuse, and other illicit substances. In response to these data gaps, this study quantifies compound-specific concentrations, examines spatial variability, and evaluates ecotoxicological risk across diverse urbanized systems. The following section outlines the sampling design, analytical protocols, and risk assessment framework used in this investigation.
Sample collection
Sampling was conducted at strategically selected sites along each river system between October 2022 and March 2023 by regionally based field teams (Fig. 1). This timeframe was deliberately chosen to coincide with (a) stable base flow conditions, (b) consistent hydrologic behavior across river systems, and (c) the absence of significant snowmelt contributions, thereby minimizing seasonal hydrologic variability. Where applicable, sampling sites were located both upstream and downstream of effluent discharge points from municipal and county wastewater treatment plants (WWTPs) and homeless encampment tents/sites.
Sampling protocol involved collecting samples from up to ten (10) locations along each river and was based on established methodologies validated in prior regional and international studies, including those by Sims et al., (2024), Wilkinson et al. (2022) and Zarębska et al. (2024). In select cases, fewer sites were sampled due to limited site accessibility, safety considerations, or adverse environmental conditions.
At each site, surface water was aspirated approximately one meter from the shoreline using a sterile syringe, and 20 mL of sample was immediately filtered through a Whatman® 0.45 µm glass fiber syringe disc filter into a 40 mL amber VOA vial. Each vial was labeled with the sampling location, date, time, and field team identifier. Samples were stored in insulated containers with pre-frozen ice packs and shipped overnight (within 24 hours) to the College of Southern Nevada laboratory in Las Vegas, Nevada (USA) for analysis. All sampling locations for each river can be viewed in Supplementary Table 1 (S1).
Chemicals and reagents
High-purity reference standards (> 98%) were sourced from a range of reputable suppliers, including Acros Organics (NJ, USA), Alfa Aesar (MA, USA), Spectrum Chemical (NJ, USA), Tokyo Chemical Industry (TCI, OR, USA), MP Biomedicals (CA, USA), American Radiolabeled Chemicals (MO, USA), Chemodex (St. Gallen, Switzerland), Cerilliant (TX, USA), Restek (PA, USA), and Cayman Chemical (MI, USA). The deuterated internal standard, carbamazepine-D10, was obtained from Santa Cruz Biotechnology (TX, USA).
All solvents used in the study were LC-MS grade. Methanol (> 99.9%) and water were purchased from Avantor Performance Materials (NJ, USA). Aqueous formic acid (> 98%) was sourced from TCI (OR, USA), while ammonium formate (1 M) was obtained from Rigaku Reagents (WA, USA). Filtration was performed using EZFlow 0.45 µm, 25 mm diameter glass fiber syringe disc filters from Foxx Life Sciences (NH, USA).
A total of twenty-eight analytes were monitored. These are atenolol, psilocin, acetaminophen, oxycodone, cefpodoxime, trimethoprim, amphetamine, MDMA, methamphetamine, ciprofloxacin, lidocaine, sulfamethoxazole, Ketamine, benzoylecgonine, fentanyl, nor-fentanyl, metoprolol tartrate, LSD, azithromycin, PCP, carbamazepine, fluoxetine, ketoprofen, bezafibrate, naproxen, diazepam, atorvastatin, and diclofenac acid.
Individual stock solutions of each compound were prepared in LC-MS grade methanol at concentrations of 1000 mg/L. A mixed working stock solution containing all target compounds was prepared at 400 µg/L in methanol and stored in 10 mL amber glass vials at -20°C in the dark to preserve stability.
LC-MS/MS analysis
Sample analysis followed a modified protocol based on methods described by Sims et al. (2024) and Wilkinson et al. (2022), in accordance with USEPA Method 1694 guidelines. Briefly, 1 mL of each water sample was transferred to a vial and placed into a temperature-controlled autosampler maintained at 4°C. From each vial, a 50 µL aliquot was directly injected into a PerkinElmer LX50 UHPLC system coupled with a PerkinElmer QSight 420 triple quadrupole mass spectrometer, equipped with a vertical mass detector and dual ionization source (ESI/APCI), operated in multiple reaction monitoring (MRM) mode.
Instrument parameters, including collision energy (CE), entrance voltage (EV), and lens voltage (CCL2), were individually optimized for each target analyte. Analyses were conducted using positive electrospray ionization (ESI+), with at least two transition ions monitored per compound to ensure both quantification and confirmation. Electrospray voltage was set at 5,500 V, with a source temperature of 320°C and Hot Surface Induced Desolvation (HSID) temperature of 280°C.
Chromatographic separation was achieved using a PerkinElmer Quasar SPP C18 column (2.6 µm, 150 × 3.0 mm) at a flow rate of 0.4 mL/min. Mobile phase A consisted of LCMS-grade water containing 0.01 M formic acid and 0.01 M ammonium formate, while mobile phase B was 100% methanol. The gradient began at 10% B, increased to 100% by 10 minutes, held until 20 minutes, then returned to 10% at 20.5 minutes, followed by a 10-minute equilibration before the next injection.
Quantification of all target analytes was performed using an 8-point external calibration curve, spanning a concentration range of 15.6 ng/L to 8000 ng/L. Calibration standards were matrix-matched to ensure accuracy and reproducibility. Each calibration curve contained a fixed concentration of isotopically labeled internal standards, selected to match the physicochemical properties and retention of a group of target compounds. Internal standards were used for response normalization, correcting for matrix effects, ion suppression/enhancement, and instrumental drift. Calibration curves were evaluated using weighted (1/x or 1/x²) linear regression, with correlation coefficients (R²) exceeding 0.995 for all analytes. Method performance was further validated through back-calculation of concentrations to confirm accuracy across the full calibration range. All instrument control, data acquisition, and processing were managed using Simplicity™ 3Q Software.
Quality assurance and quality control
All glassware was thoroughly cleaned and maintained to prevent cross-contamination. After each use, glassware was rinsed three times with a 50:50 (v/v) mixture of acetonitrile and acetone, scrubbed with LC-MS grade water, and then soaked in 10% nitric acid for a minimum of 16 hours. Volumetric flasks were designated for specific compounds to avoid cross-contamination during storage and handling. All project and QA/QC samples were analyzed in triplicate to ensure reproducibility.
To monitor instrument performance and maintain data quality during LC-MS/MS analysis, a Continuing Calibration Verification (CCV) standard and a Continuing Calibration Blank (CCB) were injected after every tenth sample, following USEPA guidelines (USEPA 1997). The CCV was prepared from a reference standard obtained from a different vendor than the calibration standards to ensure external validation, while the CCB consisted of LC-MS grade water. Each analytical batch also included a Laboratory Control Sample (LCS), prepared using a certified reference standard from a third vendor, to further confirm accuracy of quantification.
Multiple types of blanks were incorporated to evaluate background contamination, including method blanks (MB), field blanks (FB), and reagent blanks (RB). All blanks were treated identically to environmental and calibration samples in terms of storage, handling, and analysis. For example, blanks consisted of 20 mL LC-MS grade water stored in solvent-rinsed 40 mL amber glass vials at 4°C until analysis. None of the blanks, i.e. MB, FB, RB, or CCB, contained detectable concentrations of the target analytes, confirming the absence of background contamination.
To assess the effectiveness of glassware decontamination procedures, a laboratory rinse test was performed by subjecting 20 mL of a 50:50 (v/v) acetonitrile:methanol solution containing 1% acetic acid to 20 minutes of ultrasonication at 45°C, followed by LC-MS/MS analysis (n = 3). No pharmaceuticals or illicit drugs (PIDs) were detected, indicating the laboratory cleaning protocols were effective.
Each analytical batch included a Low Limit of Quantification (LLOQ) standard to confirm instrument sensitivity at the lowest calibration level. These LLOQ standards were prepared using an independent source and served as a quality check for accurate quantification near detection limits.
The LCS samples included all target PIDs at varying concentrations in LC-MS grade water, with acceptance criteria set at ± 25%, consistent with USEPA QA/QC standards (USEPA 1997). Each batch included one LCS. Internal Calibration Verification (ICV) and CCV samples were prepared at 125 ng/L with acceptance limits of ± 15% of the true value. An ICV was run at the beginning of each sequence, followed by CCVs after every 15 samples.
Matrix spike (MS) recovery checks were also included, using spiked samples at 4,000 ng/L (mid-range of the calibration curve) with acceptance windows of ± 25%. One MS and three matrix spike duplicates (MSD1, MSD2, MSD3) were analyzed after every 20 samples, following USEPA QA/QC protocol guidelines. All environmental samples, blanks, control samples, and calibration standards were spiked with 500 ng/L of carbamazepine-D10 as the internal standard to monitor extraction and injection consistency.
Occurrence of PIDs in effluent and/or unsheltered homelessness-impacted sites
This section presents the occurrence and distribution of pharmaceuticals and illicit drugs (PIDs) in surface waters across selected river systems in the Western United States, with emphasis on areas influenced by treated wastewater and unsheltered populations. The analysis includes both point sources (e.g., WWTP effluent) and non-point sources (e.g., human waste from encampments), offering a comparative assessment of contamination patterns. Results are organized by river system to highlight geographic variability in compound occurrence and concentration, and to examine potential ecological and human health implications. All data referenced in this section is available in Supplementary Table 2 (S2).
Boise River
The Boise River serves as a representative example of an effluent- and homelessness-impacted system in the Western U.S. The following results highlight both pharmaceutical diversity and environmental variability at this site. These areas combine point sources, such as treated effluent from wastewater treatment plants, with non-point sources like direct human waste inputs. Rivers such as the Los Angeles, Santa Cruz, and Rio Grande showed consistently higher PID levels than less urbanized or effluent-isolated systems, challenging assumptions about pollutant sources and treatment effectiveness.
Although many compounds, such as psilocin, oxycodone, amphetamine, MDMA, ciprofloxacin, and norfentanyl, were undetected or below quantification limits, acetaminophen was found at measurable levels. Its median concentration in the Boise River (13.2 ng L⁻¹) aligned with recent national data (Bradley et al., 2016; Henderson et al. 2020), though it remained lower than values reported in earlier studies (Kolpin et al. 2004; Bunch and Bernot 2011). Boise Creek, a critical spawning site for Chinook and coho salmon, showed detectable but unquantified acetaminophen (Tian et al. 2021), highlighting the interplay between human activity and ecological sensitivity in shaping contaminant profiles.
The human antimicrobials trimethoprim and sulfamethoxazole were detected at median concentrations of 5.41 ng L⁻¹ and 12.3 ng L⁻¹, respectively (Fig. 2). In agreement with these results, Bueno et al. (2021) reported mean trimethoprim and sulfamethoxazole concentrations of 8.6 ng L⁻¹ and 19.8 ng L⁻¹, respectively, in surface waters in Minnesota. Gray et al. (2020) reported a higher detection frequency for sulfamethoxazole (>75%) in rural streams in the Piedmont region of North Carolina. Sulfamethoxazole is widely administered with trimethoprim to treat various infections in a 5:1 ratio (Kemnic and Coleman 2023). Overall, antimicrobials—including trimethoprim and sulfamethoxazole, have been detected in U.S. surface waters at concentrations ranging from non-detectable to μg L⁻¹ levels (Kolpin et al. 2002; Gray et al. 2020). The concentrations measured in Boise River samples are several orders of magnitude below the inhibitory concentration (1 μg L⁻¹) for aquatic organisms stipulated by the United States Food and Drug Administration (1997).
Notable detections were observed for lidocaine (median = 35 ng L⁻¹) and metoprolol tartrate (median = 41.2 ng L⁻¹) in the Boise River (Fig. 2). Their presence may indicate low removal efficiency of pharmaceutical residues in local wastewater treatment plants (WWTPs). Asimakopoulos et al. (2017) previously found that lidocaine was scarcely removed by secondary wastewater treatment. Similarly, Rúa-Gómez and Püttmann (2012) indicated that lidocaine is only partially removed from sewage using conventional WWTPs. Rúa-Gómez et al. (2012) further noted that lidocaine could not be effectively removed using activated sludge treatment technologies and is therefore continuously discharged into receiving surface waters. The City of Boise’s Water Renewal Services operates two WWTPs, Lander Street and West Boise Water Renewal Facilities, both utilizing activated sludge treatment technology. Most effluent from these facilities is discharged into the Boise River, making it likely that lidocaine enters the river from these sources (Ryan 2023).
Beta-blockers are widely prescribed for treating hypertension and cardiovascular diseases. They are among the most commonly prescribed medications in the United States, with approximately 30 million adults using them. Depending on the specific drug, beta-blockers are typically dosed at least twice daily (Farzam and Jan 2023). Among β-blockers, metoprolol tartrate poses increasing concern because it is not readily biodegradable. Its rising usage and the low efficiency of WWTPs in removing this water-soluble compound contribute to its presence in waterways (Evans et al. 2015; Khan et al. 2017). In recent years, metoprolol tartrate has frequently been detected in U.S. surface waters (Minnesota Department of Health (MDH) 2015; Sims et al. 2024). Other studies have also reported average concentrations exceeding 40 ng L⁻¹ in Dagu Drainage Canal, China (Xu et al., 2019); River Fyris, Sweden (Daneshvar et al. 2010); and surface waters in Luxembourg (Singh et al. 2021). In the Boise River, the β-blocker atenolol was also detected but at a lower median concentration of 8.17 ng L⁻¹ (Fig. 2). Metoprolol tartrate has also been reported at the highest concentration among β-blockers in the BKT Tampoi catchment, Malaysia (Al-Odaini et al. 2013).
Methamphetamine use is rising in the United States. For example, usage among 12th-grade students increased by 60% from 2015 to 2017 (Kann et al. 2018), with the Pacific and Western regions particularly affected (Skeer et al. 2021). Methamphetamine was detected in 80% of Boise River water samples, with a median concentration of 8.12 ng L⁻¹ (Fig. 2). Methamphetamine has also been reported in Bee Creek (86.4 ng L⁻¹) and the Clarks River (9.47 ng L⁻¹), which flows into the Ohio River (Skees et al. 2018). Subedi and Kannan (2014) evaluated the removal efficiency of two WWTPs in New York State and found that methamphetamine showed negative removal rates, suggesting it can persist through activated sludge and biological treatment systems, such as those used in Boise. The psychiatric drug carbamazepine was detected in 60% of Boise River samples, with a median concentration of 3.48 ng L⁻¹. Skees et al. (2018) reported a much higher mean carbamazepine concentration (63.1 ng L⁻¹) in Bee Creek, a stream that directly receives treated wastewater, thus accounting for higher PID detections at that site.
Non-steroidal anti-inflammatory drugs (NSAIDs), including diclofenac and ketoprofen, are among the most frequently used pharmaceuticals in both human and veterinary medicine. Their widespread consumption has led to their ubiquitous presence in U.S. aquatic environments (Vanderford and Snyder 2006; Sims et al. 2024). In this study, diclofenac and ketoprofen were detected in 60% of Boise River samples, with median concentrations of 8.34 ng L⁻¹ and 2.99 ng L⁻¹, respectively (Fig. 2). Multiple studies have also identified these compounds in European rivers and streams (Fernández et al. 2010; Johnson et al. 2013; Vystavna et al. 2018). For example, Marsik et al. (2017) reported concentrations approaching 1 μg L⁻¹ in rivers within the Elbe basin of the Czech Republic.
While the Boise River revealed moderate contamination from select β-blockers and analgesics, more urbanized systems such as the Los Angeles River exhibited significantly higher concentrations of PIDs, likely reflecting both higher population density and more intense anthropogenic activity.
Los Angeles River
Some of the PIDs studied, including cefpodoxime, amphetamine, MDMA, ciprofloxacin, LSD, azithromycin, PCP, fentanyl, fluoxetine, bezafibrate, diazepam and atorvastatin, were either not detected or detected but not quantified. Nonsteroidal anti-inflammatory drugs (NSAID), including naproxen, ketoprofen and diclofenac acid, tend to be detected in water samples from Los Angeles River more frequently, and at relatively higher concentrations, than most other surface water systems in the U.S. These observations are similar to findings reported by Boyd et al. (2003), Lin et al. (2015), and Cabeza et al. (2012) for U.S., Taiwanese, and Spanish aquatic environments, respectively. Metoprolol succinate, metoprolol tartrate and atenolol together account for more than 63% of total β-blocker consumption in the U.S. (Definitive Healthcare 2024). The median concentrations of atenolol and metoprolol tartrate in Los Angeles River were 78 ng L-1 and 705 ng L-1, respectively (Fig.3). Previously, several studies reported metoprolol concentrations higher than 1 μg L-1 in rivers in U.S. and Europe (Ternes 2001; BLAC 2003).
Similarly, concentrations of antimicrobials such as trimethoprim, sulfamethoxazole and azithromycin measured in the Los Angeles River samples are several levels of magnitude higher than the concentrations in other river systems. These antibiotics have been frequently detected in surface water and treated wastewater in southern California (California Research Bureau (CRB) 2014). Other pharmaceuticals detected in Los Angeles River included psilocin, acetaminophen and carbamazepine with median concentrations of 15 ng L-1, 62.9 ng L-1 and 156 ng L-1, respectively (Fig.3).
Amphetamine-type stimulants, particularly MDMA and methamphetamine, enter receiving water bodies by human excretion after legal or illegal consumption via WWTPs (Boles and Wells 2010). Selected Amphetamine-type stimulants are primarily excreted as the intact drug. In the present study, amphetamine and MDMA were hardly detected in Los Angeles River. On the other hand, the median concentration of methamphetamine in water samples from Los Angeles River was 73.2 ng L-1 (Fig.3). Previous wastewater-based epidemiology (WBE) studies have reported high methamphetamine loads in wastewater from southern California cities (Pelley 2019; Watanabe et al. 2020). Notably, methamphetamine per-capita consumption rate (5.4 g/d/1000 people) determined in a study by Pelley (2019) was one of the highest rates ever reported by WBE technique for the U.S. Bishop et al. (2020) also reported remarkable levels of methamphetamine (Sit A: 0.984 g/day per 1000 people and Site B: 6.51 g/day per 1000 people) flowing through two representative western U.S. rural communities. Benzoylecgonine, the metabolite of cocaine, was also quantitated in 90% of water samples from Los Angeles River with a median concentration of 24 ng L-1. Because benzoylecgonine is a metabolite unique to human excretion, its presence suggests a chronic and continuous condition of human sewage contamination in southern California stream (Watanabe et al. 2020).
Lidocaine was detected in excessive concentration in water samples from Los Angeles River that can be used as an adulterant to ‘cut’ cocaine due to its synergistic effects (Broséus et al. 2015) or is used as local anaesthetic but its removal from WWTPs is low. Several studies have been performed to evaluate the presence and identification of cutting agents on seized cocaine exhibits collected across the globe and the active compounds detected included lidocaine and levamisole (Schneider and Meys 2011; de Oliveira Penido et al. 2016; Fiorentin et al. 2019).
It has been reported that opioids have increasingly been misused and caused public health concerns in Midwest, pacific and Western US (Gushgari et al. 2019; Havro et al. 2022). For example, the CDC estimates that over 110,000 people in the U.S. died from drug overdoses in 2022, almost 70% of these deaths were caused by fentanyl and other synthetic opioids (United States Drug Enforcement Administration (DEA) 2023). In the present study, the parent compound fentanyl and fentanyl metabolite norfentanyl were included in the analytical method. The median concentration of norfentanyl was 114 ng L-1 in water samples from Los Angeles River (Fig.3). However, fentanyl was hardly detected. Previously, Gushgari et al. (2019) reported that concentrations of norfentanyl in treated wastewater were significantly higher than (from 2-times to 48-times) the corresponding concentrations of parental fentanyl. Similarly, Fernandez (2022) found 22 positive occurrences for norfentanyl in New York city waterways, whereas only one sample was positive for fentanyl. This could be ascribed to rapid in vivo degradation and transformation of fentanyl following administration (Labroo et al. 1997; Rodayan et al. 2016; Gushgari et al. 2019).
The median concentrations of the other opioids, ketamine and oxycodone, in water samples from Los Angeles River were 45.5 and 15.2 ng L-1, respectively (Fig.3). In a recent study, Restrepo-Vieira et al. (2010) identified ketamine at very high concentrations in raw wastewaters in Australia and they also reported ketamine's persistence at WWTPs, which resulted the occurrence of ketamine in treated wastewater at concentrations as high as 212 ng L-1. Previously, Miller et al. (2019) reported higher concentrations of ketamine in surface water, the source of which was unclear. Our recent study has also revealed quantifiable concentrations of ketamine in Las Vegas wash (Sims et al., 2024). Previously, oxycodone had been detected in Bee Creek (Kentucky) at a mean concentration of 27 ng L-1 (Skees et al. 2018).
Rio Grande River
Moving into the interior southwestern U.S., the Rio Grande River provides a useful contrast between urban and peri-urban PID loadings. Samples were collected at two different sites on the Rio Grande River: Albuquerque urban area and Los Lunas. Several PIDs, including atenolol, acetaminophen, oxycodone, lidocaine, sulfamethoxazole, ketamine, norfentanyl, metoprolol tartrate, carbamazepine, and diclofenac acid were detected in significant concentrations in Albuquerque urban area compared to concentration found in Los Lunas. For example, two β-blockers, atenolol and metoprolol tartrate, tend to be detected at significantly higher concentrations in Albuquerque urban area than the concentrations found in water samples collected from Los Lunas village (Fig.4a, b). The high potential for pharmaceuticals such as β-blockers to enter river water that flows through urban areas is not unique to Rio Grande River; it has been studied in other surface water systems in the U.S. (Batt et al. 2016; Huerta et al. 2018; Bradley et al. 2020). Nonsteroidal anti-inflammatory drugs (NSAID), particularly naproxen (27.4 ng L-1), and diclofenac acid (48.7 ng L-1), occurred in far higher median concentrations in Albuquerque than concentrations found in Los Lunas. Previously, McQuillan et al. (2002) detected no NSAID in Rio Grande in Albuquerque South Valley. This was, in fact, expected because of at least two reasons; the detection limit for the targeted pharmaceuticals in their study was 10 ng L-1 and the Albuquerque population increased by nearly 24% between 2000 and 2023.
Similarly, the median concentrations of acetaminophen and carbamazepine in river water samples collected in Albuquerque were 12.2 ng L-1 and 58.1 ng L-1, respectively (Fig. 4a). Interestingly, methamphetamine was quantified with a median concentration of 5.51 ng L-1 in Los Lunas while the compound wasn’t quantified in Albuquerque urban area. On the other hand, amphetamine, benzoylecgonine and MDMA were not detected in quantifiable concentrations in both Albuquerque urban area and Los Lunas village. Sulfamethoxazole was detected in 90% of water samples collected in Albuquerque with a median concentration of 41. 2 ng L-1. Previously, Brown et al. (2006) reported way higher sulfamethoxazole concentration (300 ng L-1) in Rio Grande in Albuquerque. In fact, sulfamethoxazole appeared to be resistant to transformation and persisted in the river over approximately four miles downstream of WWTP with no change in concentration (Brown et al. 2006). Overall, the significant variability of PIDs on the Rio Grande River as it flows through Albuquerque and Los Lunas is explained by significant release associated with higher PID consumption by increasing number of urban inhabitants. Likewise, Cardini et al. (2021) reported a growing gradient of pharmaceutical concentrations from the rural to the urban areas in the coastal area of Central Italy. According to Mandaric et al. (2017) and Cardini et al. (2021), the spatial variability of pharmaceutical concentrations can be explained not only by the higher density of resident population and thus by the higher pharmaceutical consumption and release, but also by specific activities normally performed in urban areas, such as medical care and services, tourism, sport, and hospitality. Very recently, Beisner et al. (2024) assessed the temporal and spatial variability of PFAS in the Rio Grande near Albuquerque and reported an order of magnitude increase of PFAS as it flows through the Albuquerque urban area.
Santa Cruz River
In contrast to the WWTP-impacted sites along the Rio Grande, the Santa Cruz River is primarily influenced by unsheltered encampments, offering an important lens into non-point source contamination. In the U.S, unsheltered homelessness is characterized by tents or sleeping bags on urban sidewalks, and along rivers and other waterways (Flanigan and Welsh 2020). Illicit drugs are sometimes sold and shared inside temporary homeless encampment tents. In this context, rivers and other waterway shores offer a form of escape from public view and routine police inspection (Calderón-Villarreal et al. 2022). In addition to the social plight associated with unsheltered homelessness, it has potential implications for urban water quality, as wastewater generated by homeless people is discharged untreated into waterways. Other activities introducing PIDs into Arizona’s surface water system include urban stormwater runoff, and effluent discharges (Arizona Department of Environmental Quality (ADEQ) 2016). For Santa Cruz River, water samples were collected downstream of temporary homeless encampment sites. The methamphetamine concentrations in Santa Cruz River ranged from 83.5 to 450 ng L-1 (Fig.5). These high concentrations clearly reflect direct human inputs. Previous studies demonstrated similar results (Calderón-Villarreal et al. 2022; Gerrity et al. 2022). For example, Jones-Lepp et al. (2012) reported a high methamphetamine concentration (570 ng L-1) in Santa Cruz River. The median concentration of MDMA in Santa Cruz River was 10.3 ng L-1. On the other hand, amphetamine wasn’t detected in all water samples, which corroborate previous studies establishing that amphetamine is susceptible to WWTP treatments (Kasprzyk-Hordern et al. 2009; Andrés-Costa et al. 2014). Benzoylecgonine was also quantitated in 90% of water samples from Santa Cruz River with a median concentration of 8.87 ng L-1. Elevated levels of lidocaine ranging from 1306 to 2048 ng L-1 were found in Santa Cruz River (Fig.5).
The median concentrations of antibiotics ranged from 6.94 ng L-1 (Ciprofloxacin) to 626 ng L-1(sulfamethoxazole) (Fig.5). We found that trimethoprim, azithromycin and sulfamethoxazole were detected in higher concentrations. Antibiotics that are ubiquitously consumed but passes through conventional WWTPs altered slightly (e.g., trimethoprim, azithromycin and sulfamethoxazole (Karthikeyan and Meyer 2006; Bhandari et al. 2008) also dominated in Santa Cruz River, suggesting direct human contamination from homeless encampments and reduced removal in WWTPs (Gerrity et al. 2022). Several other common compounds, including β-blockers such as atenolol and metoprolol tartrate, and NSAIDs such as naproxen and diclofenac acid, were also detected in higher concentrations in Santa Cruz River relative to other water systems, further supporting the direct input theory. Similarly, psychiatric drugs carbamazepine and fluoxetine were detected in 100% and 70% of water samples from Santa Cruz River with median concentrations of 384 ng L-1 and 31.5 ng L-1, respectively. Other pharmaceuticals, including psilocin (median = 13.6 ng L-1), acetaminophen (median = 28.7 ng L-1), ketoprofen (median = 10.2 ng L-1) and bezafibrate (median = 26.6 ng L-1) were quantified in relatively lower concentrations. On the other hand, cefpodoxime, PCP, LSD, diazepam and atorvastatin weren’t detected.
Opioid misuse among youths is an increasing problem. Opioid use for homeless young people is higher in comparison to their housed counterparts (Gomez et al. 2010; Coombs et al. 2023). A very common opioid (i.e., fentanyl) has been detected in 60% of the water samples with a relatively lower median concentration value (4.83 ng L-1), whereas fentanyl metabolite norfentanyl was detected in 100% water samples with a higher median concentration of 278 ng L-1 (Fig.5). This corroborated with Maricopa County (Arizona) media advisory noting an increase in local fentanyl deaths by 4900% since 2015 (Marcopa County 2022). The median concentrations of the other opioids, ketamine and oxycodone, in water samples collected from Santa Cruz River were 29.4 and 66.5 ng L-1, respectively (Fig.5).
Truckee River
The Truckee River, which receives treated effluent but has fewer encampment-related inputs, showed a narrower range of detected PIDs. Examples of pharmaceuticals that weren’t detected or quantified in Truckee River include psilocin, cefpodoxime, amphetamine, MDMA. PCP, fentanyl, carbamazepine, ketoprofen, bezafibrate, naproxen, and diazepam. Similarly, antibiotics have been hardly detected in water samples collected from Truckee River. On the other hand, reclaimed water discharged from Truckee Meadows Wastewater Reclamation Facility (TMWRF) into the Truckee River contained high levels of sulfamethoxazole (~ 500 ng L-1) and trimethoprim (~ 1500 ng L-1) (Sharma et al. 2020). Some TMWRF reclaimed water is used to irrigate local golf courses and the University of Nevada, Reno Experiment Station’s Main Station Field Laboratory (MSFL). A field-scale study at MSFL demonstrated that carbamazepine and trimethoprim were translocated to shoots and leaves of alfalfa irrigated with reclaimed water (Sharma et al., 2020).
β-blockers are of concern as they are the most frequently detected pharmaceutical classes in aqueous systems; atenolol and metoprolol were detected in 100% of US surface waters (Zhang et al. 2021). In the present study, the median concentrations of quantified β-blockers in Truckee River varied from 18.2 ng L-1 (atenolol) to 97.7 ng L-1 (metoprolol tartrate). Because of their frequent use, β-blockers have been widely detected in surface waters in Europe and North America with concentrations ranging from a few ng L−1 up to 2.2 μg L−1 (Ternes 2001; Zhang et al. 2021). In these settings, higher concentrations were measured in surface waters impacted by municipal wastewaters (Alder et al. 2010; Sims et al. 2024).
An analgesic drug acetaminophen was detected in 60% of water samples from Truckee River with a median concentration of 7.16 ng L-1. A systematic review conducted by Zhang et al. (2021) regarding a vast array of contaminants of emerging concern in water cycle in the US demonstrated that carbamazepine levels in surface water ranged from 0.01 ng L-1 and 10 μg L-1. Anticonvulsants drug carbamazepine was detected in 70% of water samples from Truckee River with a median concentration of 18.0 ng L-1. Previously, Park and Park (2015) reported that the mean carbamazepine level in Lake Mead, located in the southwestern region between the states of Nevada and Arizona, was 30.7 ng L-1. Nonsteroidal anti-inflammatory drugs except diclofenac acid were hardly detected in Truckee River.
Anesthetic (lidocaine and ketamine) concentrations in Truckee River ranged from <LOD to 13.6 ng L-1 (Fig.6). The median concentrations of opioid (i.e., oxycodone) and metabolite of fentanyl (i.e., norfentanyl) in water samples from Truckee River were 6.5 ng L-1 and 17.3 ng L-1, respectively (Fig.6). Zhang et al. (2021) demonstrated that oxycodone levels in US surface water ranged from 0.01 ng L-1 and 100 ng L-1. Similarly, a recent pharmaceuticals and personal care products monitoring study by Gerrity et al. (2024) in Southern Nevada sewershed (combined storm and sewer systems) revealed a stark increase in fentanyl consumption beginning in October 2022. Methamphetamine was detected in 70% of the samples with a median concentration of 4.94 ng L-1(EDC). For benzoylecgonine, median concentration was 10.4 ng L-1 (Fig.6). This is consistent with our recent study that revealed relatively lower concentrations of benzoylecgonine in Las Vegas wash (Sims et al. 2024).
Virgin River
To round out this regional assessment, the Virgin River represents a less urbanized system in a desert watershed, enabling comparisons across urban–rural and humid–arid gradients. Located in the southwest United States where Utah, Nevada, and Arizona converge is where the Virgin River is located. It is a tributary of the Colorado River and provides essential water resources for millions of people for drinking, irrigation and recreation. In the present study, water samples were collected from the Virgin River within Utah. Concentrations of detected antibiotics varied by more than 3 orders of magnitude from less than 1 ng L−1 to more than 4 μg L−1. In the last decade, many reviewers have compiled data on the distribution of antibiotics in the U.S. surface waters (Deo and Halden 2013: Fang et al. 2019; Wilkinson et al. 2022). For example, Fang et al. (2019) had critically reviewed pharmaceuticals including antibiotics in surface waters from 25 states in the United States. They found that the concentrations of antibiotics in surface waters ranged from 5 ng L-1 to 1500 ng L-1.
Beta blockers are one of the most used classes of drugs in the U.S. The median concentration of atenolol in East River in New York city (Wilkinson et al. 2022) was approximately double the median concentration obtained in the Virgin River. On the other hand, higher median concentration (199 ng L-1) was exhibited for metoprolol tartrate in the Virgin River compared to atenolol (Fig.7). When comparing metoprolol tartrate and atenolol prescriptions dispensed in the U.S. and dosage forms, metoprolol tartrate prescribed highly, twice as much as atenolol, and tends to have high recommended dosing frequency and maximum total daily dose. Anesthetics, such as lidocaine and ketamine, median concentrations ranged from ND to 399 ng L-1 in the Virgin River (Fig.7). Moreover, the median concentrations of opioid (i.e., oxycodone) and metabolite of fentanyl (i.e., norfentanyl) ranged between 20.5 ng L-1 and 6.2 ng L-1, respectively (Fig.7).
Non-steroidal anti-inflammatory drugs (NSAIDs), such as naproxen and diclofenac, were detected in 70% of water samples with median concentrations detected at 26.5 ng L-1 and 91 ng L-1, respectively (Fig.7). Several studies have also detected NSAIDs in higher concentrations in U.S. surface waters (Boyd et al. 2003; Vanderford and Snyder 2006; Wu et al. 2009). Similarly, the median concentrations of psilocin and carbamazepine in the Virgin River were 11.4 ng L-1 and 60.4 ng L-1, respectively. In the present study, amphetamine, MDMA, methamphetamine and benzoylecgonine were hardly detected. These results align with Jones-Lepp et al. (2012) who reported no detection for methamphetamine and MDMA in the Virgin River. Other non-detected or non-quantified PIDs included acetaminophen, cefpodoxime, LSD, MDMA, PCP, fentanyl, fluoxetine, ketoprofen, bezafibrate, diazepam and atorvastatin.
Overall, sites influenced by wastewater effluent and unsheltered homelessness exhibited elevated concentrations of specific pharmaceutical and illicit drugs (PIDs), indicating complex contamination dynamics. These environments reflect both point sources, such as effluent from WWTPs, and non-point sources, including direct human waste inputs from unsheltered populations. The resulting PID signatures suggest sustained, multi-source contamination that challenges conventional assumptions about pollutant origin and removal efficiency in urban surface waters. Detailed analysis of water samples from the Los Angeles River, Santa Cruz River, and Rio Grande illustrates these patterns, with several compounds detected at concentrations significantly higher than those observed in less urbanized or effluent-isolated systems.
Ecotoxicological risk assessment
While spatial variability highlights differing sources and concentrations of PIDs, understanding their ecological implications requires a systematic risk assessment. To this end, both acute and chronic ecological risks were estimated using the RQ framework outlined below.
Potential ecological risks posed by specific contaminants to aquatic ecosystems were assessed following the technical guidance documents on environmental risk assessment issued by the European Commission (European Commission (EC) 2003; European Medicines Agency 2006). The risk quotient (RQ) for each contaminant was calculated as the ratio of the measured environmental concentration (MEC) to the predicted no-effect concentration (PNEC) (Eqn. 1). Risk was classified based on standard thresholds: RQ < 0.1 indicated low or negligible risk; 0.1 ≤ RQ < 1 denoted medium risk; and RQ ≥ 1 indicated high risk (Hernando et al. 2006; Hoang et al. 2024; Zhang et al. 2024). For RQmax calculations, the maximum recorded MEC of each pharmaceutical was used to assess potential acute risk under worst-case exposure scenarios. In contrast, the median MEC was used to calculate RQmedian, offering a more representative estimate of chronic or long-term exposure effects (Palma et al. 2020). PNEC values were derived from the lowest available half maximal effective concentration (EC50) or no observed effect concentration (NOEC) values for algae (S. subspicatus and P. subcapitata), crustaceans (D. magna), and fish (D. rerio), divided by an appropriate assessment factor (AF) (Eqn. 2). Toxicity data were obtained from both the ECOSAR v2.0 predictive model and peer-reviewed literature sources. Specific toxicological endpoints and concentration-response data for each compound in relation to the target species are presented in Table S3.
Both RQmax and RQmedian were utilized to provide a comprehensive evaluation of environmental risk, capturing both acute worst-case scenarios and more realistic chronic exposure conditions. This dual approach enabled a more nuanced understanding of the ecological risks associated with the target pharmaceutical compounds (Table S3; Figs.8 & 9). Results indicated that the RQmax values for acetaminophen, diclofenac, naproxen, carbamazepine, atenolol, trimethoprim, lidocaine, fluoxetine, and bezafibrate in algae were consistently below 0.1 across all effluent-impacted river systems, suggesting negligible to low ecological risk to algal species (Fig.8). These findings align with previous studies conducted in Europe and North America (Sanderson et al. 2003), as well as global assessments of pharmaceutical contamination in surface waters (Bouzas‐Monroy et al. 2022). Similarly, Liu et al. (2024) reported RQ values below 0.01 for these compounds in effluent-impacted rivers in eastern China, further supporting a negligible risk classification for algae.
In contrast, ciprofloxacin, ketoprofen, azithromycin, and metoprolol exhibited moderate ecological risks to algae. Sulfamethoxazole posed a high risk, with an RQmax value of 1.76 (Fig.8). This result is consistent with Bouzas‐Monroy et al. (2022), who found that sulfamethoxazole was frequently detected at concentrations exceeding PNEC values in rivers across Africa, Asia, and North America, with RQmax values exceeding 1.0. In North American river systems specifically, the RQmax for sulfamethoxazole reached 1.8.
Daphnia magna is a widely used and well-established invertebrate model organism in ecotoxicology and plays a critical role in aquatic food webs by serving as an intermediary between primary producers and higher trophic levels (Flaherty & Dodson 2005). In the present study, the calculated RQmax values for all analyzed pharmaceuticals, except fluoxetine, were below 0.1 across all effluent-impacted river systems (Fig. 2), indicating low or negligible ecological risk to crustaceans. These findings are consistent with previous studies conducted in North America, which have also reported low or negligible risk of pharmaceutical contaminants to D. magna, with RQmax values frequently falling below 0.01 for selected compounds (Skees et al. 2018; Bouzas‐Monroy et al. 2022; Wilkinson et al. 2022).
In contrast to algae and crustaceans, fish exhibited a broader range of susceptibility to pharmaceutical exposure in effluent-impacted river systems, with RQmax values indicating low to high ecological risk for select compounds. This heightened vulnerability is partly attributable to elevated environmental concentrations of metoprolol and the low predicted NOEC estimates for fluoxetine.
Generally, these risk quotients were derived by comparing maximum MECs, representing worst-case exposure scenarios, with the most conservative ECOSAR-estimated NOEC, EC50, or LC50 values. As a result, these RQs likely overestimate actual environmental risk (Sanderson et al. 2003). Nonetheless, relying solely on single-compound risk assessments may underestimate cumulative ecological risks, as such methods fail to account for mixture effects and potential synergistic or additive toxicity of pharmaceutical mixtures, conditions that more accurately reflect real-world environmental exposures (Wolfe et al. 2015; Bouzas‐Monroy et al. 2022).
Despite several published works on PIDs in the aquatic environment in the US, there remain significant gaps, including little monitoring of PID occurrence and distribution in urban watershed sites that are influenced by both effluent and unsheltered homeless encampments. An analytical methodology based on LC-MS/MS was utilized to determine PIDs in anthropogenically impacted sites in select US river systems, namely, Boise, Los Angeles, Rio Grande, Santa Cruz, Truckee, and Virgin. Although PID concentration in the river systems is site-specific, lidocaine was detected in excessive concentration in several effluent-receiving river systems. It can be used as an adulterant to ‘cut’ cocaine due to its synergistic effects or as a local anaesthetic, but its removal from WWTPs is low. Other pharmaceuticals and metabolites detected in higher concentrations in the Los Angeles River included atenolol, norfentanyl, carbamazepine, and metoprolol tartrate with median concentrations of 78 ng L⁻¹, 156 ng L⁻¹, 114 ng L⁻¹, and 705 ng L⁻¹, respectively. WWTPs are relevant sources of PIDs in urban watersheds, and PID concentrations might not decline over a longer river length. Likewise, the widespread occurrence of PIDs in unsheltered homelessness-impacted river systems clearly illustrates that potential environmental impact is one of the negative societal consequences of unsheltered homelessness. For example, methamphetamine concentrations in the Santa Cruz River ranged from 83.5 to 450 ng L⁻¹. Similarly, the median concentrations of antibiotics ranged from 6.94 ng L⁻¹ (ciprofloxacin) to 626 ng L⁻¹ (sulfamethoxazole).
Overall, sites influenced by wastewater effluent and unsheltered homelessness exhibited elevated concentrations of specific PIDs, indicating complex contamination dynamics. These environments reflect both point sources, such as effluent from WWTPs, and non-point sources, including direct human waste inputs from unsheltered populations. The resulting PID signatures suggest sustained, multi-source contamination that challenges conventional assumptions about pollutant origin and removal efficiency in urban surface waters. Detailed analysis of water samples from the Los Angeles River, Santa Cruz River, and Rio Grande illustrates these patterns, with several compounds detected at concentrations significantly higher than those observed in less urbanized or effluent-isolated systems. Thus, measures to mitigate unsheltered homelessness extend beyond employment assistance, low-income housing, mental health, and other social services.
Potential ecological risks posed by specific pharmaceuticals to aquatic ecosystems were also assessed using environmental risk assessment tools. However, ecotoxicity data of illicit drugs and their metabolites/transformation products is deficient. This hampers a reliable environmental risk assessment of this group of compounds. Our findings revealed that pharmaceuticals, except sulfamethoxazole, were detected at concentrations far below PNEC values across all effluent- and unsheltered homelessness-impacted river systems, indicating negligible (low) or moderate ecological risks to algae and crustaceans.
In contrast to algae and crustaceans, fish exhibited a broader range of susceptibility to pharmaceutical exposure in effluent-impacted river systems, with RQmax values indicating low to high ecological risk for select compounds. This heightened vulnerability is partly attributable to the most conservative ECOSAR-estimated NOEC, EC50, or LC50 values. For this reason, these RQs are likely to overestimate actual environmental risk. The RQmax values reported here are therefore important as they allow the identification of the ranges of maximum pharmaceutical concentrations that can normally be present in river systems without posing any risk to aquatic ecosystems. Nonetheless, relying solely on single-compound risk assessment may underestimate cumulative ecological risks, as such methods fail to account for mixture effects and potential synergy or additive toxicity of pharmaceutical mixtures.
Additionally, this study did not evaluate the potential for bioaccumulation or biomagnification of pharmaceutical compounds across trophic levels in aquatic ecosystems. While risk quotient (RQ)–based assessments provide valuable insight into species-specific toxicity thresholds, they do not capture the ecological consequences of trophic transfer or contaminant persistence in higher-order organisms. Several compounds identified in this study, including fluoxetine and carbamazepine, are known to bioaccumulate in aquatic biota and may pose long-term ecological risks via food web exposure pathways. Future research should incorporate tissue residue analysis, bioconcentration or bioaccumulation factors (BCFs/BAFs), and trophic transfer modeling to enable more holistic evaluations of pharmaceutical impacts in effluent-influenced ecosystems.
Finanly, this study underscores the urgent need to broaden environmental risk frameworks to include both chemical monitoring and social infrastructure factors, particularly the environmental consequences of untreated human waste in unsheltered settings. Effective management of urban water quality will require not only improved WWTP technologies but also integrated social policy solutions to address homelessness as an environmental stressor.
Acknowledgements
We thank the College of Southern Nevada for their continued support of student research.
Author Contributions
All authors contributed to the study conception and design. All authors performed material preparation, data collection, and analysis. All authors prepared the frst draft of the manuscript and read and approved the fnal version.
Funding
This research was partially supported by a grant from the National Institute of General Medical Sciences (GM103440) and the State of Nevada Office of Science, Innovation & Technology (OSIT).
Data Availability
All data analyzed during this study are included in the article and available through first primary author.
Conflict of interest
The authors declare they have no competing interests.
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